Exposure to commonly-used phthalates and the associated health risks in indoor environment of urban China

Science of the Total Environment 658 (2019) 843–853

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Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Exposure to commonly-used phthalates and the associated health risks in indoor environment of urban China Zhongming Bu a,⁎, Daniel Mmereki b, Jiahui Wang c, Cong Dong a a b c

Department of Energy and Environmental System Engineering, Zhejiang University of Science and Technology, Hangzhou 310023, China National Centre for International Research of Low-carbon and Green Buildings, Chongqing University, Chongqing 400045, China Institute of Urban Construction, Hangzhou Polytechnic, Hangzhou 311402, China

H I G H L I G H T S

G R A P H I C A L

A B S T R A C T

• Measurements of five phthalates in indoor environment of China were summarized. • Exposures and health risks were assessed for urban Chinese populations. • DiBP and DnBP were the highest contributors to non-carcinogenic risks. • Higher risks were from residential exposures for both children and adults. • DEHP exposure resulted in higher lifetime cancer risk for working adults.

a r t i c l e

i n f o

Article history: Received 1 September 2018 Received in revised form 21 November 2018 Accepted 17 December 2018 Available online 19 December 2018 Editor: Wei Huang Keywords: Phthalate Exposure Health risk Indoor Urban China

⁎ Corresponding author. E-mail address: [email protected] (Z. Bu).

https://doi.org/10.1016/j.scitotenv.2018.12.260 0048-9697/© 2018 Elsevier B.V. All rights reserved.

a b s t r a c t Rapid urbanization and modernization have increased exposures to phthalates from synthetic materials used indoors in China. However, exposure to phthalates from indoor environment and the associated health risks to the urban population have not been adequately characterized and documented. In this study, we summarized the recent measurements of five commonly-used phthalates in indoor environment in urban China and documented their distributions. Based on the activity patterns and exposure factors of Chinese population, Monte-Carlo simulation was used to derive their exposures. On average, the daily intake of all the targeted phthalates was 3.6 μg/kg/day for adults; and for children it ranged from 4.4 μg/kg/day to 8.1 μg/kg/day. For children, the total risk from exposures inside residences and offices was 32%–90% and 4%–19%, respectively. From commuting environments and other indoor environments, it was 5%–31%, and 3%–26%, respectively. For adults, the total risk from residences and offices was 26%–78% and 9%–35%. Additionally, from commuting environments and other indoor environments, it was 8%–35% and 5%–11%, respectively. The non-carcinogenic risk assessment was based on a cumulative Tolerable Daily Intake (TDIcum), with means ranging from 0.18 to 0.41, which was mainly as a result of exposure to DiBP and DnBP. The means for lifetime cancer risk resulting from DEHP exposure ranged from 0.4 × 10−6 to 2.0 × 10−6 for urban population groups. For 80% of working adults and 40%–75%% of children, their cancer risks exceeded the EPA's benchmark (1.0 × 10−6). The present study could provide important information for decision makers to reduce indoor phthalate exposures as well as the associated health risks for larger population groups in Chinese cities. © 2018 Elsevier B.V. All rights reserved.

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1. Introduction Phthalates, known as plasticizers, have been used to enhance the flexibility and durability of materials for decades. In modern world, phthalates are used in a wide range of daily products, including soft polyvinyl chloride (PVC), children's toys, food packages, medical products and other synthetic materials (Gao et al., 2018; Giovanoulis et al., 2018; Gong et al., 2016). Some lower-molecular-weight phthalates are also used as solvents or carriers of cosmetics and personal care products (Bu et al., 2018; Gong et al., 2015). Phthalates can be continuously released into indoor environment over time since they are not chemically bound to the products. As a consequence, phthalates are presently ubiquitous indoors and humans are exposed to them on a daily basis. Studies have shown that exposure to phthalates may cause numerous adverse health effects, including reproductive system dysfunction (Chin et al., 2018; Radke et al., 2018; Smarr et al., 2018; Toft et al., 2012), endocrine disorders (Araki et al., 2017; Arbuckle et al., 2018; Yu et al., 2018), asthma and allergies in children (Ait Bamai et al., 2018; Bornehag et al., 2004; Li et al., 2017; Shi et al., 2018), neurodevelopment problems in children (Braun, 2017; Lee et al., 2018) and obesity (Amin et al., 2018; Xia et al., 2018). Moreover, di(2-ethylhexyl) phthalate (DEHP) has been classified as a probable human carcinogen (EPA, 1988). The continuous use of phthalates in plasticized products results in a growing trend of indoor pollution from phthalates (Weschler, 2009). In the past few decades, the consumption of plasticizers has increased rapidly due to rapid urbanization and modernization in China. In 2006, phthalate consumption was approximately 1.7 million tons in China, accounting for a quarter of that consumed globally (Tao and Liang, 2008; Wang et al., 2010); and the volume increased to 2.5 million in 2011 (Qian and Zhu, 2011). Previous studies pointed out that indoor phthalate concentrations in several cities in China were at a higher level compared to those measured in other countries (Bu et al., 2016; Wang et al., 2014). Based on the recognition of the hazardous effects of phthalates, the European Union and the US issued regulations to restrict the usage of phthalates in daily products since the end of the last century (EPA, 2012; EU, 1999, 2007). Comparatively, China's national standard, which regulated the usage of phthalates in toys (GB6675-2014) firstly came into effect in 2014 (GB6675-2014, 2014). Therefore, due to the massive consumer market and the relatively recent regulatory actions, phthalate pollution in indoor environment in China is a major concern. Research interests of indoor phthalate pollution in China have increased rapidly since 2010. Exposures via multiple pathways have been quantified through environmental concentration measurements (Bu et al., 2016; Guo and Kannan, 2011; Wang et al., 2017; Wang et al., 2014; Zhang et al., 2014) or biomonitoring (Gong et al., 2015; Guo et al., 2011; Han et al., 2014). However, indoor population exposure to phthalates in urban China has not been adequately characterized, including the associated health risks. Individual studies have been done in various regions across the whole country, with localized cases using relatively small sample sizes. Most studies only focused on simpler indoor environment, such as residences (Guo and Kannan, 2011; Pei et al., 2013; Zhang et al., 2013), offices (Song et al., 2015; Wang et al., 2014), schools (Li et al., 2016), public places (He et al., 2016; Wang et al., 2017) and hospitals (Wang et al., 2015). However, they could not give in-depth and complete picture of population exposures. To the best of our knowledge, no in-depth and systematic analyses of indoor exposure and the associated health risks posed by commonly-used phthalates in multiple microenvironments have been done in China. In this study, we first summarized indoor phthalate measurements in urban China, focusing on commonly-used compounds in different microenvironments. Thereafter, personal exposure models via multipathways were derived for Chinese population based on their activity patterns indoors. The associated non-carcinogenic risk assessment was undertaken based on the European Food Safety Authority's Tolerable Daily Intake (TDI), and apportioned among different compounds as well as various indoor microenvironments. Cancer risk as a

consequence of DEHP exposure was further estimated using the proposed method by California Office of Environmental Health and Hazard Assessment (OEHHA) for carcinogens. The results could provide valuable information for decision makers to regulate the phthalate industry and reduce indoor exposures in urban China. 2. Materials and methods 2.1. Literature review and data collection A review of existing data was carried out from published literature reporting measurements of indoor phthalates in cities in China using “Web of Science” (WoS) and “China National Knowledge Infrastructure” for articles published prior to Oct., 2018. The articles were obtained using keywords or keyword sets, “phthalate or PAE”, “indoor or home” and “China”, to request topic search (including fuzzy search) by the WoS search system. The words were translated into Chinese when scoping searches for relevant published literature from the Chinese database. Most of the studies were published during the last 10 years. In the present study, investigations that used biomonitoring for exposure measurements were excluded since environmental concentrations were required for further calculations. Review articles were also excluded. The selected 20 measurements in multiple microenvironments from 14 cities presented useful information on recent phthalate pollution and characterization in urban China. As noted by Du et al. (2014), the targeted microenvironments used to determine indoor exposures included indoor residence, indoor working place (office), commuting environments and others in our study. To the best of our knowledge, investigations on commuting environments were limited to two studies with available data of airborne phthalate levels in automobiles. Considering the fact that Chinese working adults spend 60% of their commuting time in automobiles (NBS, 2010), we treated the corresponding data as representative of the commuting environments in urban China. The miscellaneous “others” included microenvironments such as dormitories, schools, shopping malls and hospitals. The selected measurements in specific cities covered five important regions in China, i.e., the Pearl River Delta (PRD) Region, the Yangtze River Delta (YRD) Region, the Bohai Rim (BR) Region, the Western Region, and the Northeast Region. Table 1 presents the selected references with detailed information on the study location, sampling media, sample size and type of microenvironments. Based on our data collection, five compounds with higher concentrations in both indoor air and settled dust were considered as the most commonly-used phthalates in urban China, i.e., dimethyl phthalate (DMP), diethyl phthalate (DEP), di(isobutyl) phthalate (DiBP), di(nbutyl) phthalate (DnBP) and di(2-ethylhexyl) phthalate (DEHP). Other phthalates were excluded from the present study due to their lower concentrations or less occurrence in reviewed articles, such as butyl benzyl phthalate (BBzP), di(n-octyl) phthalate (DNOP) and divyclohexyl phthalate (DCHP). In most of the studies, the concentrations of individual compounds in different samples were not available. Thus, the reported statistics were used for our analysis. To properly weigh statistics from distinct studies, we used the method suggested by Loh et al. (2007). Each city located in the same geographic region was equally weighted. If there was more than one study for a city, each study was usually weighted according to the unique number of sampling sites. To a lesser extent, the weighting was performed based on the number of collected samples. Afterwards, each region was equally weighted to derive the concentration distributions for urban China. 2.2. Exposure models and risk characterization Personal exposure level is one of the most important parameters for exposure assessment. Indoor exposure to phthalates occurs in numerous microenvironments. In the present study, a time-weighted

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Table 1 Recent publications reporting indoor phthalate measurements in urban China. Reference Pearl River Delta Region Lan et al. (2012) Wang et al. (2013) Guo and Kannan (2011) Wang et al. (2013) Yangtze River Delta Region Zhang et al. (2013) He et al. (2016) Wang et al. (2012) Guo and Kannan (2011) Zhang et al. (2016) Pei et al. (2013) Song et al. (2015) Wang et al. (2015) Pei (2013) Huang (2014) Bohai Rim Region Wang et al. (2017) Guo and Kannan (2011) Lin et al. (2009) Guo and Kannan (2011) Li et al. (2016) Ji et al. (2014) Zhang et al. (2014) Zhang (2016)

Location

Residence

Office

Guangzhou

√ √ √ √

Hong Kong

Nanjing

√ √

Commute

Others

Sample size

√

Apartment

10/15/10 20 11 20

√ √ √ √

√

Public place/dormitory

215 6/12/7/8 72 21 8 10 10 5 10/10 8/3/24

√ √

√ Shanghai Hangzhou

√ √ √ √

√

√

Hospital Dormitory Dormitory

√ √ √ √

√

Kindergarten/public place

√

Dormitory

√

Beijing

Ji'nan Baoding Tianjin

Dormitory √ √ √

Western Region Bu et al. (2016) Wang et al. (2014)

Chongqing Xi'an

√ √

Northeast Region Guo and Kannan (2011) Li et al. (2016) Li et al. (2016)

Qiqihaer Harbin Shenyang

√ √

Dormitory Dormitory

exposure model that integrates the time fraction spent in each microenvironment and the concentrations of each microenvironment (Dodson et al., 2007) was applied to predict personal exposure levels among the Chinese population, as shown in Eq. (1): C w;i ¼

X

C i; j t j

Dust

√ √ √ √ √ √

√ √ √ √ √ √ √ √ √ √ √ √ √

√ √

√

√ √ √ √ √

approach. A relative cumulative TDI (TDIcum) introduced by Koch et al. (2011) was applied, as shown in Eq. (2): TDI cum ¼

ð1Þ

where Cw,i is the time-weighted personal exposure level to pollutant i, μg/m3 (subscript “w” means “weighted”.); Ci,j is the concentration of pollutant i in microenvironment j, μg/m3; tj is the time fraction spent in microenvironment j. Humans are exposed to phthalates via multiple pathways, including inhalation, oral ingestion, and dermal absorption (Salthammer et al., 2018; Weschler and Nazaroff, 2008). In our study, nondietary ingestion was only accounted for oral ingestion since data on phthalate levels in food were not collected. Equations for exposure estimates via multi-pathways are listed in Table A.1 in the Supporting Information (SI). Based on the selected investigations, indoor dust-phase concentrations were major concerns. For a complete assessment of all pathways, gas-phase was estimated based on the partitioning model of semi-volatile organic compounds (SVOC) between the gas and settled dust (Weschler and Nazaroff, 2010) if only the dust-phase levels were available. For a more accurate estimation of the airborne concentrations, an “indoor particle age” concept (Cao et al., 2018a) was applied to describe the adsorption of gas-phase SVOC in indoor particles. Detailed information is shown in Section A.2 in the SI. The exposure to several endocrine active phthalates may act in a dose-additive manner. The associated non-carcinogenic health risks were assessed based on the concept of cumulative risk assessment

12 35/18 8

Particle

√

23/30/45/15 11 10/10/10 13 8 26 82 410

30 14/14

√

Gas

X DI i TDI i

ð2Þ

where DIi is the total daily intake of phthalate i, μg/kg/day; TDIi is the corresponding tolerable levels, μg/kg/day. A TDIcum greater than one indicates that the cumulative daily phthalate intake potentially leads to adverse health effects. Considering the similar health endpoints, the TDI cum was calculated based on the intakes of DiBP, DnBP and DEHP. In this study, the modeled population in China were divided into six age groups, i.e., b1 year, 1–5 years, 5–12 years, 12–18 years, 18–64 years and N 64 years. For working adults (18–64 years), activity distributions of both males and females were extracted from the Time Use Patterns in China (NBS, 2010). For other age groups, the time use patterns were based on the U.S. Environmental Protection Agency's (EPA) exposure factor handbook (EPA, 2011). For children b1 year old, our estimates were based on calculating one year's residential exposure. Exposure factors of target population groups were obtained from the Exposure Factors Handbook of Chinese Population (Ministry of Environmental Protection, 2013). Detailed distributions of each input factors are listed in Tables A.3–A.4. Among the targeted phthalates, DEHP is a probable human carcinogen (EPA, 1988). For each age interval, cancer risk was assessed based on the estimation of lifetime cancer risk (LCR), as follows: LCR ¼

X

 CPF j  CDI j  ADAF

ð3Þ

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where CPFj is the cancer potency factor for pathway j, (μg/kg/day)−1; ADAF is the age-dependent adjustment factor, which was introduced to take into account the susceptibility of exposure during the early life stage. U.S. EPA recommended an ADAF of 10 for 0–2-year-old, 3 for 2–16-year-old, and 1 for 16–70-year old (EPA, 2005). CDIj is the chronic daily intake, μg/kg/day, which is expressed as Eq. (4): CDI ¼

DI  F  ED AT

ð4Þ

where DI is the daily intake, μg/kg/day; F is the exposure frequency, day/ year, given as 365 days/year in our calculations; ED is the exposure duration, year, which is equal to the yearly duration of each age interval (Zhu et al., 2019); AT is the average lifetime (70 years), i.e., 25,500 days. For DEHP, the cancer potency values were only available for inhalation and oral intake based on OEHHA (2009). Therefore, the total cancer risk was calculated as the sum of the LCR via inhalation and non-dietary ingestion. Based on the EPA guidelines, a value of b10−6 was considered a negligible cancer risk (Caldwell et al., 1998). The TDI values and cancer potency information for targeted phthalates are presented in Table A.5. 2.3. Uncertainty and sensitivity analysis A stochastic risk assessment approach was used to deal with the uncertainty and variability of each input parameter. Monte-Carlo simulation was applied to characterize the ranges of exposures and risks. Sensitivity analysis ranks input assumptions according to their importance to the outcome. In the present study, Monte-Carlo simulation and sensitivity analysis were carried out using Oracle Crystal Ball software for Windows (Fusion Edition, 64-bit, V. 11.1.2.2). A lognormal distribution was used to describe the cumulative frequency distributions of phthalate concentrations. A coefficient of variance (CV), expressed as the ratio of the standard deviation to the mean, was introduced to describe the distributions if the reported statistics of certain parameters were incomplete. Statistical data were insufficient in five studies, where standard deviations were taken as 20% of the means to compile distributions (i.e., CV = 0.2). Exposure parameters were assumed to be lognormally distributed (CV = 0.3) except for the body weight, which was assigned as a normal distribution with a CV of 0.1 or 0.2 (Bu et al., 2018). As for physical properties of phthalates, such as partitioning coefficient and transdermal coefficient, a CV of one was assigned (Gaspar et al., 2014). For cancer potency factors

(CPF), triangle distribution was chosen to characterize the uncertainty (Zhou et al., 2011). The CPF values recommended by OEHHA were assumed to represent the maximum and the most likely values, and 0 was assigned to be the minimum. Specific values of physical properties are listed in Table A.6. Thereafter, 50,000 trials of Monte-Carlo simulation (i.e., enough to reach a stable mean or standard deviation) were performed to calculate the exposures and risks. 3. Results and discussion 3.1. Commonly-used phthalates in urban China Most studies focused on the direct measurement of indoor dustphase concentrations. Consequently, based on the weighted statistics of the targeted regions, the concentrations of the three most predominant phthalates in indoor dust across urban China were obtained. Based on (Du et al., 2014), a full map of China with respect to the national dust-phase levels is presented in Fig. 1. For each microenvironment, the contribution of each phthalate was similar. DEHP was most predominant in settled dust, followed by DnBP and DiBP. In urban China, dust-phase DEHP concentrations varied by a factor of 20, from b100 μg/g to N2000 μg/g, but the typical weighted medians ranged from about 500 μg/g to 900 μg/g. The levels of the two isomers, DiBP and DnBP, were in order of magnitude lower than DEHP. For other two phthalates, i.e., DMP and DEP, the dust-phase levels were much lower (medians b10 μg/g) due to a lower partitioning coefficient between dust and air. The concentrations from commuting environments were not shown since the data from direct measurements were not available. Statistical data extracted from each study were detailed in Table A.7. For the three most predominant phthalates in settled dust, a comparison was made among the different geographical regions for the dust-phase levels since the levels of DMP and DEP were much lower. Our comparison focused on the residential microenvironment since the relevant data were sufficient to present all targeted regions across urban China. As shown in Fig. 2, the dust-phase levels were comparable among the targeted regions except for the Western Region, where the measured phthalates in settled dust were roughly 3–10 times higher than other regions. The relevant data were obtained from investigations in Chongqing (Bu et al., 2016) and Xi'an (Wang et al., 2014), respectively. The measured levels of DiBP and DnBP in Xi'an were roughly 1000 μg/g and 750 μg/g, respectively, which were obviously higher than the average values across urban China. A possible explanation

Fig. 1. Distributions of measured dust-phase phthalate concentrations across urban China.

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than that for urban China (147 μg/g). For indoor air concentrations (including the particle-phase), the levels of DMP (1.1 μg/m3) and DnBP (1.2 μg/m3) were slightly higher than those reported by our study (0.73 and 0.87 μg/m3, respectively), while the levels of DEP (0.54 μg/ m3) and DEHP (0.30 μg/m3) were lower than our findings (0.82 and 1.8 μg/m3, respectively). The DiBP level was roughly 8 times lower than that for urban China. A recent study focusing on integrated exposure to phthalates in China also provided comparable findings to our study. Cao et al. (2016) summarized indoor phthalates concentrations based on field data from six mega-cities in China. The indoor dust levels of DiBP, DnBP and DEHP were 15 μg/g, 25 μg/g and 243 μg/g, respectively (geometrical means), which were 4–10 times lower than the reported levels in our study. The levels of DiBP, DnBP and DEHP in indoor air were more comparable with our study. 3.2. Modeled exposures

Fig. 2. Dust-phase phthalate levels in residences from different geographical regions.

could be a relatively shorter period of occupancy after construction (b5 years) for the selected buildings. For Chongqing city, the measured DEHP level were roughly 3 times higher than the average level. However, the concentrations of urinary metabolites for Chongqing adults (Han et al., 2014; Liu et al., 2012) were somewhat lower than those of the general population in China (Guo et al., 2011). A comparative study regarding large scale investigations of indoor environment, and also human biomonitoring is needed to provide an in-depth and comprehensive analysis of the inconsistencies. Table 2 lists the gas-phase concentrations of the most commonlyused phthalates derived from available studies. Gas-phase concentrations in commuting environments were higher than other microenvironments. DiBP concentrations were highest in working environments. Although DiBP levels were absent for commuting environments, the sum of all phthalate concentrations was the highest (3.9 μg/m3), suggesting that the air in automobiles was significantly polluted. The interior materials were the major sources of airborne SVOC in automobiles, including phthalates (Hoshino et al., 2005) and polybrominated diphenyl ethers (Mandalakis et al., 2008). Notably, the concentrations of SVOC declined with vehicle's age and increased with the rise in temperature. The total gas-phase concentrations for offices, homes and “others” were 3.7, 2.5 and 2.0 μg/m3, respectively. To the best of our knowledge, previous studies reporting on nationwide indoor phthalate concentrations did not distinguish between the different microenvironments. Thus, in our study, we averaged the mean concentrations from different microenvironments for easier comparison. Wormuth et al. (2006) summarized measured phthalate concentrations in different media in European countries. For the dustphase, the distribution patterns of each compound were similar to those observed in our study in urban China. To be specific, DEHP concentration (1198 μg/g) was somewhat higher than reported (874 μg/ g) in our study, while DiBP concentration (84 μg/g) was slightly lower

Table 2 Statistical summary of gas-phase concentrations in microenvironments (μg/m3). Phthalates

DMP DEP DiBP DnBP DEHP

Arithmetic mean (standard deviation) Residence (n = 985)

Office (n = 94)

Commute (n = 82)

Others (n = 168)

0.82 (1.21) 0.38 (1.80) 0.71 (1.67) 0.54 (0.74) 0.09 (0.04)

0.90 (1.37) 0.98 (2.02) 1.23 (3.59) 0.45 (0.64) 0.09 (0.05)

1.10 (0.50) 1.30 (0.31) – 0.82 (0.63) 0.71 (0.58)

0.64 (0.76) 0.61 (0.67) 0.29 (0.68) 0.35 (0.79) 0.11 (0.05)

Tables 3–4 shows statistical summary of the modeled exposures for typical population groups in urban China. For children, the daily intakes of larger molecular weight phthalates (DiBP, DnBP and DEHP) were higher than for the other two phthalates. The total exposure of DEHP was the highest, followed by DiBP and DnBP. The inhalation exposure was comparable for each phthalate. The non-dietary ingestion of DEHP was the highest, while the ingested doses were in order of magnitude lower for DMP and DEP due to their lower dust-phase concentrations. For DiBP and DnBP, the amount of dermal absorption was higher than the other phthalates. For adults, the proportional distribution of each pathway was similar to that of children. The exposure levels were similar for both males and females. Non-dietary ingestion was the most predominant pathway for DEHP intake, contributing approximately 60% and 70% for adults and children, respectively. Dermal absorption for the most part contributed the daily intakes of DEP, DiBP and DnBP (roughly 40%–70%). The total exposures were higher for children's groups than adults' groups. The daily intakes of younger group (b1 year) were even higher, which indicated a potential health risk for the infants. We compared our exposure estimates with results from previous studies, including estimations based on indirect approach via multipathways as well as biomonitoring. As listed in Table 5, estimated exposure levels (medians) varied among the different population groups in different countries. For adults' exposures, the levels ranged from 0.23–1.15 μg/kg/day for DEP, 0.2–1.7 μg/kg/day for DiBP, 0.34–8.7 μg/kg/day for DnBP and 0.16-11 μg/kg/day for DEHP. Our results showed lower levels of exposure, generally lower than those estimated through biomonitoring or food intakes. Similar observations were made for children's exposures, indicating lower estimates. A possible explanation could be that exposures from other microenvironments were not effectively quantified in our study, such as the sleeping microenvironments, daycare centers and kindergartens. Noticeably, DEHP exposure was roughly 3 times higher than that for children in Denmark, when similar pathways as our study were considered. 3.3. Health risk assessment Fig. 3 presents the estimated cumulative TDI (TDIcum) based on the daily intakes of the targeted phthalates with similar health endpoints. The cumulative TDI for males (18–64 years) was shown as a representative since similar results were observed for adults' groups (both males and females), with a median value of 0.13. The results for children were possibly higher than those for adults, with medians from 0.15 to 0.30. The cumulative TDI decreases with age, with the highest value for infants b1 year. For N95% of the adult population, the cumulative TDI was less than one. The proportion of children with a TDIcum less than one accounted for about 90%. The results suggested acceptable non-carcinogenic risks resulting from phthalates exposures for Chinese populations.

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Table 3 Modeled children's exposure via multiple pathways (μg/kg/day)a. Pathways

Arithmetic mean (standard deviation) DMP

DEP

DiBP

DnBP

DEHP

b1 year Inhalation Non-dietary ingestion Dermal absorption Total

0.46 (0.80) 0.01 (0.01) 0.23 (0.60) 0.69 (1.24)

0.21 (1.07) 0.01 (0.02) 0.35 (1.80) 0.58 (2.60)

0.45 (1.03) 0.48 (0.81) 0.88 (3.00) 1.81 (3.81)

0.37 (0.53) 0.52 (0.49) 0.71 (1.68) 1.60 (2.10)

0.57 (0.92) 2.67 (1.98) 0.15 (0.20) 3.39 (2.29)

1–5 years Inhalation Non-dietary ingestion Dermal absorption Total

0.40 (0.53) – 0.23 (0.52) 0.62 (0.89)

0.23 (0.80) 0.01 (0.01) 0.40 (1.86) 0.64 (2.34)

0.37 (0.84) 0.27 (0.42) 0.81 (2.93) 1.47 (3.50)

0.32 (0.36) 0.30 (0.24) 0.67 (1.41) 1.30 (1.91)

0.55 (0.70) 1.55 (0.94) 0.18 (0.27) 2.28 (1.25)

5–12 years Inhalation Non-dietary ingestion Dermal absorption Total

0.31 (0.38) – 0.22 (0.47) 0.53 (0.73)

0.20 (0.48) 0.01 (0.01) 0.45 (1.66) 0.67 (3.39)

0.29 (0.62) 0.28 (0.42) 0.79 (2.59) 1.38 (4.23)

0.25 (0.26) 0.30 (0.23) 0.64 (1.18) 1.20 (1.66)

0.49 (0.52) 1.67 (0.98) 0.19 (0.27) 2.35 (1.27)

12–18 years Inhalation Non-dietary ingestion Dermal absorption Total

0.22 (0.27) – 0.19 (0.40) 0.41 (0.53)

0.15 (0.35) 0.01 (0.01) 0.40 (1.55) 0.55 (1.51)

0.21 (0.42) 0.16 (0.23) 0.68 (2.25) 1.04 (2.48)

0.17 (0.17) 0.16 (0.13) 0.65 (1.06) 0.98 (1.27)

0.36 (0.38) 0.92 (0.55) 0.16 (0.24) 1.43 (1.79)

a

The values lower than 0.005 were treated as missing data.

We ranked the health risks by making a comparison of the mean risks of individual compound. Although the health risks varied among the different age groups, the differences did not alter the ranking of the pollutants. Taking the males for an example (shown in Fig. 4), DiBP and DnBP accounted for the highest portion of the total risk, contributing roughly 50% and 40%, respectively. DEHP accounted for a lower portion, contributing 11%. The contributions of each microenvironment to the total risk were compared based on the calculated TDIcum for different populations (Fig. 5). For children, on average, the total risk from the exposures inside residences and offices was 32%–90% and 4%–19%, respectively. From the commuting environments and other indoor environments, it was 5%– 31% and 3%–26%, respectively. For the adults, 26%–78% of the total risk

was from residences, 9%–35% from offices, 8%–35% from commuting environments, and 5%–11% from other indoor environments. It was worth noting that the residential contributions decreased with age for children, whereas those contributions increased with age for adults. Our findings were consistent with the fact that the younger and the elderly tend to spend more time at home. The health risk from residential exposure was slightly higher for working females, while the risk from exposure in offices was to some extent higher for working males. It should be born in mind that the exposures from commuting environments was comparable to or even higher than the risks from offices. Therefore, the air quality in commuting environments needs to be given more consideration when making policies and decisions concerning pollution reduction.

Table 4 Modeled adults' exposure via multiple pathways (μg/kg/day)a. Pathways

Arithmetic mean (standard deviation) DMP

DEP

DiBP

DnBP

DEHP

18–64 years (male) Inhalation Non-dietary ingestion Dermal absorption Total

0.22 (0.24) – 0.15 (0.29) 0.37 (0.47)

0.16 (0.33) – 0.34 (1.08) 0.51 (1.36)

0.22 (0.36) 0.10 (0.14) 0.58 (1.69) 0.89 (2.09)

0.18 (0.16) 0.10 (0.06) 0.43 (0.76) 0.71 (0.91)

0.35 (0.30) 0.55 (0.26) 0.14 (0.19) 1.04 (0.48)

18–64 years (female) Inhalation Non-dietary ingestion Dermal absorption Total

0.21 (0.23) – 0.16 (0.34) 0.37 (0.49)

0.14 (0.32) – 0.34 (1.09) 0.50 (2.11)

0.20 (0.36) 0.12 (0.17) 0.59 (1.66) 0.91 (2.08)

0.17 (0.17) 0.12 (0.08) 0.46 (0.86) 0.75 (0.98)

0.31 (0.28) 0.66 (0.32) 0.15 (0.19) 1.12 (0.51)

N64 years (male) Inhalation Non-dietary ingestion Dermal absorption Total

0.20 (0.24) – 0.16 (0.33) 0.36 (0.55)

0.12 (0.32) – 0.33 (1.30) 0.46 (1.66)

0.19 (0.37) 0.11 (0.16) 0.62 (1.81) 0.92 (2.22)

0.16 (0.17) 0.11 (0.08) 0.50 (1.02) 0.76 (1.06)

0.28 (0.29) 0.62 (0.32) 0.14 (0.19) 1.04 (0.53)

N64 years (female) Inhalation Non-dietary ingestion Dermal absorption Total

0.20 (0.24) – 0.17 (0.37) 0.37 (0.53)

0.12 (0.29) – 0.35 (2.46) 0.46 (1.23)

0.19 (0.36) 0.12 (0.17) 0.63 (1.80) 0.97 (2.88)

0.16 (0.15) 0.13 (0.09) 0.51 (1.05) 0.81 (1.27)

0.29 (0.29) 0.68 (0.36) 0.15 (0.21) 1.13 (0.57)

a

The values lower than 0.005 were treated as missing data.

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849

Table 5 A comparison of phthalate exposures (μg/kg/day) reported in previous studies. References Estimates for adults Wormuth et al. (2006)a Fromme et al. (2007)b Fromme et al. (2004)c Giovanoulis et al. (2018)a Clark et al. (2011)a Christensen et al. (2014)b Cao et al. (2016)a Guo et al. (2011) Present studyd

b

Estimates for children Wormuth et al. (2006)a Fromme et al. (2004)c Bekö et al. (2013)b Bekö et al. (2013)d Clark et al. (2011)a Christensen et al. (2014)b Guo et al. (2011)b Present studyd a b c d

Country

Subjects

DEP

Europe

Male

1.15 –

3.61

2.85

Germany Adult Germany Adult Norway Adult

– 1.7 0.23 – 0.94 0.68

1.7 0.34 0.56

2.2 0.16 0.84

USA USA

Adult Male

0.46 0.76 – 0.2

1.2 0.5

11 3.9

China

General population Male Male (18–64 years)

–

1.00

3.02

3.80

1.0 – 0.26 0.44

8.7 0.47

3.0 0.95

Europe

Child

–

–

1.21

1.78

Germany Denmark Denmark USA USA

Child 3–6 years 3–6 years 5–11 years 6–11 years

0.61 – 0.80 0.93 –

– 2.93 1.95 2.1 0.4

0.82 3.26 0.97 2.4 0.9

6.03 4.42 0.83 20 6.0

China China

b10 years 5–12 years

0.8 – 0.30 0.70

6.1 0.81

3.4 2.06

China China

DiBP DnBP DEHP

Fig. 4. Mean contributions to total risk from different phthalates.

3.4. Sensitivity analysis Although the contributions varied among the age groups for children or adults, the differences did not alter the ranking of the parameters. Sensitivity data based on the “5–12 years” and “18-64 years (males)” age groups were chosen as representatives for children and adults, respectively. The contributions of major parameters to the total

Exposure via multi-pathways including food ingestion. Estimated through biomonitoring approach. Exposure via inhalation and dust ingestion. Estimated via inhalation, non-dietary ingestion and dermal absorption.

Fig. 6 presents the lifetime cancer risk of typical population groups as a consequence of DEHP exposure. It has been estimated that the average life expectancy for Chinese population was 74 and 79 years old for males and females, respectively (NBS, 2017). The mean values for children ranged from 1.0 × 10−6 to 1.8 × 10−6, with the highest value being for the “5–12 years” group. The mean cancer risks were approximately 2.0 × 10−6 for the working adults. For the elderly, the risks were lower, with means ranging from 0.4 × 10−6 to 0.6 × 10−6. It could be found that the lifetime cancer risk for working adults was significantly higher than other population groups. Based on the EPA's guideline value (10−6), nearly 80% of the working adults in urban China had a potential cancer risk resulting from DEHP exposure, while 40%–75% of the children may be exposed to higher cancer risk.

Fig. 3. Cumulative frequency distribution of TDIcum for populations in urban China.

Fig. 5. Mean contributions to the total risk from each microenvironment. (a) Children. The suffixes 1–3 represent the “1–5”, “5–12” and “12–18” years old, respectively. (b) Adults. The suffixes 1–3 represent the “18–64 (males)”, “18–64 (females)” and “N64” years old, respectively.

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Fig. 7. Contributions of major parameters to the total variance in TDIcum for typical populations. Sensitivity data based on “5–12 years” and “18–64 years (males)” age groups were chosen as representatives for children and adults, respectively.

Fig. 6. Population cancer risk resulting from DEHP exposure in urban China. (a) Children; (b) Adults.

variance in cumulative TDI for typical groups are presented in Fig. 7. For adults, the results of sensitivity analysis showed variations in gas-phase concentrations of DiBP and DnBP in residence, time fraction spent in home, and transdermal permeabilities of DiBP and DnBP being the predominant contributors to the total variance (roughly 70% of the total). For children, gas-phase DiBP and DnBP had the highest contributions. Other major contributors were body weight, time fraction in home and transdermal permeabilities. The estimates on which parameter is mostly influencing the risk estimates, suggested the best way to refine the estimates when more detailed data are available. 3.5. Limitations for the present work and suggested future study There was insufficient measurement data on non-home microenvironments for the targeted phthalates, especially in commuting environments, contributing to major difficulties in deriving exposures. Only two studies reported airborne concentrations in commuting environments; but the DiBP concentrations were not clear. Also, phthalate concentrations in workplaces were derived from field data from five cities, leading to uncertainty on whether the relevant data could represent the real picture of phthalate exposure from offices in urban China. Moreover, the problem of limited data still arose when we tried to compile the concentration data for the “others” microenvironments. There were few

studies reporting on phthalate concentrations in public places such as hospitals, shopping malls, schools or restaurants, while the measurements were much more for dormitories. Thus, phthalate levels from dormitories were used as surrogates for part of “others” microenvironments. In addition, the investigations were all limited and localized cases; consisting of relatively smaller sample sizes. On the other hand, only five phthalates were included in our study based on the relevant literature searches. Field test data regarding the indoor pollution from the emerging phthalates and phthalate alternatives, such as diisononylphthalate (DINP), diisodecyl phthalate (DIDP), di-2ethylhexyladipate (DEHA), di(2-ethylhexyl) terephthalate (DEHT) and 1,2-cyclohexane dicarboxylic acid diisononyl ester (DINCH), were somewhat limited. Future exposure assessment could, however, be greatly facilitated by national or regional scale measurements focusing on more compounds with sufficient sample size covering a larger population. The paucity of direct measurement of airborne concentrations also resulted in considerable uncertainty for exposure estimates. In most cases, the distributions of gas-phase concentrations were inferred from the measured dust-phase concentrations based on a steady-state model (Weschler and Nazaroff, 2010); and the particle-phase levels were further calculated from the inferred gas-phase levels. These estimates could lead to considerable errors for phthalates with a larger octanol-air partitioning coefficient (Koa), for instance, DEHP, which requires much more time to achieve equilibrium between different phases (Liu et al., 2013; Weschler and Nazaroff, 2008). The gas-phase concentrations could be underestimated from the dust-phase since the settled dust could hardly saturate in real environment, so did the particle-phase. On the other hand, indoor airborne concentrations of phthalates were also influenced by other factors such as temperature, air exchange rate (AER), and particle concentration, which could be hardly described in one study. Although the estimation of particlephase concentrations was improved by applying the “indoor particle age” concept (Cao et al., 2018a), the relevant input parameters (e.g., AER, particle concentrations) for the model were assumed based on previous studies. We acknowledged that more field tests with regards to airborne concentrations could provide more helpful information for future study. With the exception of the mentioned microenvironments in the present study, other microenvironments, including sleeping microenvironments, bar, gym, garage, etc., were not considered. Based on the previous studies, a variety of chemicals (e.g., VOCs, phthalates) have been

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detected from bedding materials (Boor et al., 2014; Boor et al., 2017; Liang and Xu, 2014), which could exist as a continuous emission source in the sleeping microenvironments. As a result, the concentrations of air pollutants could be higher in the breath zone than the bulk air when people are using the mattress during sleeping time (Boor et al., 2014; Liang and Xu, 2014), resulting in higher exposure levels, especially for infants or young children. Dermal exposure via contact with the bedding materials or even the air gap between skin and the material (Cao et al., 2018b) could also be considerably higher since the mass contents of phthalates in the crib mattress cover could be as high as 10% (Liang and Xu, 2014). However, we could not effectively quantify the exposures at this stage due to the limited data for Chinese residences. Therefore, further in-depth research focusing on the field test in those microenvironments (e.g., airborne levels, temperature, air flow rates), and also the emission rates of typical source materials is needed. The present work focused on exposures resulting from contact with environmental indoor media in China. However, time use patterns in microenvironments for pupils and the elderly were based on the available data for Americans (EPA, 2011) since detailed information was limited for the Chinese population. As for the pathways, previous studies concluded that food was important, especially for larger-molecularweight phthalates (Cao et al., 2016; Clark et al., 2011; Wormuth et al., 2006). The exposure pathways considered in the present study, however, excluded the dietary ingestion, which inevitably underestimated the total phthalate exposure for the urban population. Moreover, when calculating inhaled doses, the particle deposition in human lung (Deng et al., 2018; Hofmann, 2011; Salma et al., 2015) or the desorption of particle-phase SVOC in the respiratory tract (Liu et al., 2017) was not considered, which might result in errors of intakes via inhalation. For dermal exposure estimates, a fraction of total skin exposed to air was assumed, which disregarded the effect of clothing. Morrison et al. (2016) pointed out that wearing clothing could either accelerate or impede the absorption of gas-phase SVOC. In addition, the influence of particles on mass transfer of phthalates between air and skin surface lipids were not considered (Cao et al., 2017). Based on our results, the abovementioned exposure scenarios were somewhat not given much attention, which were limited by the insufficient data or analytical approach at this stage. A biomonitoring approach could be applied in further studies to improve the estimation of phthalate exposure and risk assessment. These are areas for future refinements. 4. Conclusions This study summarized recent measurements of five commonlyused phthalates in indoor environment and documented their distributions in urban China. The analysis of indoor phthalate concentrations from five geographical regions in urban China showed that DiBP, DnBP and DEHP were the three major phthalates. The residential exposure was the major contributor among the different microenvironments, followed by the exposure from offices and commuting environments. The cumulative TDI were mainly contributed by DiBP and DnBP, and generally less than one for most population. The cancer risk from DEHP exposure exceeded the EPA's guideline value (10−6) for 80% of the adults and 40%–75%% of children in urban China. The present study provides valuable information for commonlyused phthalate exposures and the associated health risks in indoor environment in urban China. This can be helpful to decision makers to develop control strategies and regulatory actions to reduce phthalate exposures in China. Further research on national scale measurements, especially for airborne concentrations, refined exposure assessment via multiple pathways and detailed exposure factors for children is required. Declaration of interest The authors declare no competing financial interest.

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Acknowledgements This work was supported by the National Natural Science Foundation of China (Grant No. 51606169), Natural Science Foundation of Zhejiang Province (Grant No. LQ19E080002) and the Research Foundation of Hangzhou Polytechnic, China (Grant No. HKZYYB-2017-1). Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi. org/10.1016/j.scitotenv.2018.12.260. References Ait Bamai, Y., Araki, A., Nomura, T., Kawai, T., Tsuboi, T., Kobayashi, S., Miyashita, C., Takeda, M., Shimizu, H., Kishi, R., 2018. Association of filaggrin gene mutations and childhood eczema and wheeze with phthalates and phosphorus flame retardants in house dust: the Hokkaido study on environment and children's health. Environ. Int. 121 (Pt 1), 102–110. Amin, M.M., Ebrahimpour, K., Parastar, S., Shoshtari-Yeganeh, B., Hashemi, M., Mansourian, M., Poursafa, P., Fallah, Z., Rafiei, N., Kelishadi, R., 2018. 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